Groundwater pollution by halogenated, and particularly chlorinated solvents is a worldwide problem associated primarily with industrial sites where mishandling or improper disposal has brought these solvents in contact with the soil. The most common and problematic compounds are the chlorinated ethylenes such as tetra- tri- or di-chloroethylene. Carbon tetrachloride, chloroform and methylene chloride are also pervasive pollutants. The reasons for concern are basically threefold. First, most of these solvents are sparingly soluble in water and have the tendency to stick to soil particles. This results in tenacious underground plumes of solvent which cannot readily be removed by standard pump and treat technology (Biswas, N., et al., Water Environ. Res. 64, 170, 10, 1 (1992); Hutter, G. M., et. al., Water Environ. Res. 64, 69, (1992)). Second, the toxicology of many chlorinated solvents suggests that these compounds may be carcinogenic and damaging to specific organs such as the liver and kidneys (Price, P. S., Memo of the U.S. Environmental Protection Agency, Office of Water, Washington, D.C.(1985); Vogel, T. M., Environ. Sci. Technol., 21, 722, (1987)). Finally, under conditions found in many aquifers and subsurface environments, chlorinated ethylenes and methanes are very slow to be degraded biologically. The result of these factors is that chlorinated solvents are long-lived potentially hazardous groundwater pollutants.
Currently there are two approaches to in situ removal of organohalogen pollutants. The first approach is the standard "pump and treat" method where groundwater is pumped to the surface for physical stripping of the contaminant from the water. For chlorinated solvents this is more of a containment method than a remediation technology although given sufficient time (typically decades to centuries) this method may capture most of the pollutant. The other approach is biological in nature and utilizes microorganisms for the enzymatic transformation of the halogenated organics. The biological approach may utilize microorganisms indigenous to a particular site where the remediation process consists primarily of making additions to the contaminated site that enhance the growth of the desired microorganism. Alternatively, nonindigenous microorganisms may be introduced to a contaminated site with the necessary amendments needed for growth. Dispersal of introduced microorganisms due to the filtration effect of aquifer sediments appears to be a major drawback to this approach.
A very limited number of pure strains of bacteria have been described which dehalogenate chlorinated solvents. Colaruotolo et al. (U.S. Pat. No. 4,511,657) claim the use of specially adapted microbial cultures to treat obnoxious waste, especially halogenated organic chemical waste (U.S. Pat. No. 4,493,895). EP 461144 teaches the use of members of the genera Rhodococcus and Mycobacterium to remediate chlorophenolic compounds from polluted soil and U.S. Pat. No. 5,009,999 disclose a biodegredation process for polychlorinated bi-phenyl compounds utilizing a strain of Pseudomonas putida. Additionally, strains of Klebsiella oxytoca have been demonstrated to degrade chlorobenzoic acid in liquid medium and use mono and di-chlorobenzoates as sole carbon sources (U.S. Pat. No. 4,761,376). Studies with crude anaerobic bacterial consortia have demonstrated both in the laboratory and in situ that chlorinated ethylenes or methanes can be completely dechlorinated to ethylene and methane respectively. (Freedman, D. L., et al., Appl. Environ. Microbiol., 55, 2144, (1989)). This transformation occurs via a series of less chlorinated intermediates in successive reductive steps, however, the organism or organisms responsible have not been readily purified from these consortia.
Certain species of the anaerobic sulfate-reducing and methanogenic bacteria appear to have an enhanced ability to dehalogenate halogenated solvents when compared to other bacteria. However, this trait appears to be rare even among these two groups. Desulfomonile tiedjeii, a unique isolate, is one of the most active dechlorinating anaerobes described to date (Mohn, W. W., et al., Microbiological Reviews, 56, 482, (1992)). This organism has been shown to dechlorinate tetrachloroethylene, but is more efficient against chloroaromatics. Desulfobacter sp, methanogenic bacteria and certain acetogenic bacteria such as Acetobaterium woodii have also been shown to dechlorinate chloro-aliphatics at very low rates (Egli, C. S., et al., FEMS Microbiol. Lett., 68, 207 (1990)). Egli et al. teach that when these bacteria are incubated under anaerobic conditions in reduced buffer, suspensions of each organism degraded carbon tetrachloride forming less highly chlorinated methanes and carbon dioxide as products. The physiological significance of this activity to the organism is unknown and may simply be a fortuitous sidereaction with little energetic or nutritional consequence to the organism. The microbiology of anaerobic dehalogenation of chlorinated solvents is only currently being resolved and is largely phenomenological. The biochemistry of anaerobic dehalogenation is poorly understood being limited by the extraordinarily low rates of dehalogenation exhibited by pure cultures (Mohn, W. W., et al., Microbiological Reviews, 56, 482 (1992)).
The methods cited above are useful and clearly show that microorganisms can be used to degrade halogenated organics from both soil and aqueous environments. There are, however, several disadvantages to the methods outlined in the existing art. Examples given in the art describe decontamination of the environment using specific naturally occurring, or genetically engineered cultures of bacteria or the preliminary harsh chemical treatment of toxic contaminants prior to biological treatment with microbes. It should be noted that chemical treatment of the contaminated area is likely to kill the majority of the indigenous microbial flora and thus impede any effort at bioremediation. The isolation or engineering, culturing and inoculation of specific microorganisms particularly selected for the degradation of specific organic contaminants is labor intensive and time consuming. Furthermore, as previously mentioned, in the absence of oxygen, degradation of chlorinated hydrocarbons proceeds very slowly. Years of costly groundwater recirculation may be required for conventional bioremediation to completely mineralize a pollutant. Bioremediation is very poorly understood in terms of the taxonomic types of organisms required for dechlorination. It is commonly observed that a bioremediation strategy successful at one contaminated site may not work at another. Furthermore the microbiology of some sediments cannot be induced to dechlorinate pollutants under any conditions. Certain pollutants such as perchloroethylene, for example, may be biologically transformed to vinyl chloride which is far more toxic than the original parent compound. Thus every attempt at bioremediation must be preceded by costly laboratory microcosm studies and pilot field studies to determine if dechlorinating organisms are present and, if so, what nutritional supplements need to be added to the sediments to encourage growth of these organisms.
Abiotic anaerobic dehalogenation of chlorinated hydrocarbons by iron-, nickel-, or cobalt-containing tetrapyrroles has been reported by several investigators. Krone, U. E., et al., Biochemistry, 28, 4908 (1989) teach the use of corrinoids to catalyze the reductive dehalogenation of carbon tetrachloride in conjunction with either titanium(III) citrate or dithiothreitol (DTT) as electron donors. The degradation products formed by this process included chloroform, methylene chloride, chloromethane, and CH.sub.4, indicating that complete dechlorination is possible. Additionally, Krone, U. E., et al. (Biochemistry, 28, 10061 (1989)) , demonstrated that corrin and corrinoid catalyzed dechlorination of carbon tetrachloride may also be mediated by the nickle-containing porphinoid, coenzyme F430 found in methanogenic bacteria. Marks T., et al. (WO 8910772) teach a method for the dehalogenation of organohalogen compounds by reacting the organohalogen with a reducing agent in the presence of a selected metal-centered porphyrins, corrins or phthalocyanine complexes. The method of Marks et al. encompasses the use of a complex comprising the above mentioned ring structures in association with metals selected from groups 2, 5, 6, 8, 9, or 10 of the Periodic Table.
Gantzer, C. J., et al., (Environ. Sci. Technol., 25, 715 (1991)) demonstrate that the bacterial transition-metal coenzymes vitamin B12 (containing cobalt), coenzyme F430 (containing nickle) and hematin (containing iron) catalyzed the reductive dechlorination of polychlorinated ethylenes and benzenes. Gantzer discloses that for vitamin B12 and coenzyme F430 the reductive dechlorination rates for different classes of perchlorinated compounds had the following order: carbon tetrachloride&gt;tetrachloroethylene&gt;hexachlorobenzene. For hematin, the order of reductive dechlorination rates was carbon tetrachloride&gt;hexachlorobenzene&gt;tetrachloroethylene. Within each class of compounds, rates of dechlorination decreased with decreasing chlorine content. In the reductive dechlorination of trichloroethylene, cis-1,2-dichloroethylene was the predominant product formed with vitamin B12, coenzyme F430, and hematin. Pentachlorobenzene and pentachlorophenol were each dechlorinated by vitamin B12 to yield two out of three possible isomeric tetrachlorobenzenes.
The abiotic methods involving tetrapyrroles disclosed in the art are useful for the dehalogenation of organic solvents however the requirement for the presence of a strong reductant limits their use in a subsurface application where the concentration and chemical nature of the required reductant would be prohibitive from both a regulatory and economic perspective. Thus, tetrapyrrole catalysis has, until now, only been considered an ex situ technology.
One factor essential to the metabolism of Desulfovibrio, which is potentially capable of acting as a strong reductant, is the periplasmic (external to the cytoplasmic membrane) electron carrier protein, cytochrome c.sub.3. This protein is localized on the exterior of the cell and is thus accessible to external electron transfer reactions. The electron transfer proteins from Desulfovibrio appear to be very diverse and contain at least four different c type cytochromes, including the monohemic cytochrome c553 (Mr 9000), the tetrahemic cytochrome c.sub.3 (Mr 13000), the octahemic cytochrome c3 (Mr 26000) and a high molecular weight cytochrome (Mr65000) called Hmc, containing sixteen hemes (Haladijian, J., et al., Biochem. Biophys. Res. Com., 179, 605 (1991)). A common pattern of the cytochrome c3 superfamily is that they contain c-type hemes having low redox potentials within the approximate range of -120 to -400 mV. The cytochrome c.sub.3 proteins function within the cell to transport electrons between hydrogenase and the rest of the electron transfer chain. To date, the reducing power of the periplasmic cytochrome c.sub.3 has not been coupled to any process for the dehalogenation of compounds.
From the art it is clear that a need remains for a method of dehalogenating halogenated organic pollutants in situ reliably and reproducibly at degradation rates higher than currently possible in biological systems. It is the object of the present invention to meet such a need by providing a method that utilizes indigenous populations of sulfate-reducing bacteria in combination with tetrapyrrole catalysts to effect the dehalogenation of halogenated organics in situ. The instant method utilizes sulfate reducing bacteria with the associated c.sub.3 cytochrome to provide electrons to a tetrapyrrole (vitamin B12) to catalyze the dechlorination of various chlorinated solvents. The present invention is an improvement over the art in several ways. The present method utilizes bacteria indigenous to most aquifer systems to provide the strong reductant to drive tetrapyrrole catalysis. The site need not contain dechlorinating organisms, only sulfate-reducing bacteria are required. Secondly, operationally this process would be the same regardless of the site, i.e., generation of nutritional conditions for sulfate-reducers and once they are established addition of the dechlorinating reagent. Thus, the remediation would resemble a process rather than a costly R&D project aimed at determining if dechlorinating bacteria are present.
Additionally, the instant method should provide a dramatic increase in rate of bioremediation and the overall reliability and reproducibility of the practice of bioremediation. The present invention requires only that the inherent microbiology of any sediment contain sulfate-reducing bacteria. Numerous field and laboratory studies have shown these organisms to be generally present in sediments and soils such that their presence can be relied upon (Vogel, T. M., Environ. Sci. Technol., 21, No. E., pp 77-81 (1987)); (Mohn, W. W., et al., Microbiological Reviews, 56, 482 (1992)); Suflita, J. M., et al., J. Ind. Microbiol., 3,179, (1988)). The nutritional characteristics of the sulfate-reducing bacteria are well resolved so that nutritional strategies to raise this group of bacteria are straightforward. Indigenous dechlorinating bacteria are not required but should not be inhibited by this process.